Measurement of the different forms of zinc in ... - Science Direct

Marine Chemistry, 33 ( 1991 ) 171-186

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Elsevier Science Publishers B.V., Amsterdam

Measurement of the different forms of zinc in Narragansett Bay water based on the rate of uptake by a chelating resin Francois L.L. Muller* and Dana R. Kester~ Graduate School of Oceanography, Universityof Rhode Island, Narragansett, R102882 (U.S.A..) (Received July 5, 1988; revision accepted August 30, 1990)

ABSTRACT Muller, F.L.L. and Kester, D.R., 1991. Measurement of the different forms of zinc in Narragansett Bay water based on the rate of uptake by a chelating resin. Mar. Chem., 33:171-186. Batch treatment with Chelex-100 ion exchange resin was used to quantify the physico-chemical forms of zinc in Narragansett Bay water samples collected on seven dates between November 1982 and January 1986. Labile complexes of zinc in solution were rapidly taken up ( 100% uptake after 0.5 h); another two kinetically distinguishable components of the zinc pool were identified as 'moderately labile' organic complexes and 'slowly labile' particulate forms, respectively. The first component followed pseudo-first-order kinetics of dissociation, whereas the particulate component gave a removal plot that could best be described as zero order. Although the distribution of zinc between these components varied considerably between sampling dates, the rate constant characterizing the release of zinc from each component showed little variation with season and tidal phase (dissolved organic complex, kd = 0.36 + 0.08 h - i; particulate form, kd = 0.12 + 0.04 nM h - ~). One sample differed from the above analysis in that the removal of particulate zinc was found to conform to a surface binding site kinetic analysis, from which the first-order desorption rate constant (kd=0.043 h - l ) was obtained. The environmental implications of these findings are discussed.

INTRODUCTION

In an estuarine system, trace metals are likely to interact significantly with both dissolved organic compounds and the suspended load of particles. The implications of such interactions for marine ecology were considered by Hirose and Sugimura ( 1985 ), using the concept of metal-buffering capacity. According to these authors, the role of naturally occurring organic ligands in seawater is to keep the free metal ion concentration nearly constant, in the range required for phytoplankton growth. For a system with one metal and *Present address: Department of Oceanography, The University, Southampton SO9 5NH, Gt. Britain. **Author to whom correspondence should be addressed.

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F.L.L. MULLER AND D.R. KESTER

one ligand, the magnitude of this buffering capacity and its variation with increased or decreased metal concentration can be calculated if the conditional stability constant and the organic ligand and metal concentrations are known. Furthermore, Hirose and Sugimura argued that constancy of the buffering capacity is essential for the sustained growth of phytoplanlcton and that a drop in buffering capacity as a result of metal uptake by phytoplankton may bring about the end of a bloom. This concept was applied by Hirose and Sugimura ( 1985 ), using copper as the essential 'nanonutrient', to predict the permissible period of blooms in the marine environment (4-15 days). When making such an assessment of the role of metal-organic interactions in the marine environment, it is assumed implicitly that the reactions (association/ dissociation) are reversible and sufficiently fast that equilibria among the species are maintained during the bloom period. The suggestion that this might not always be the case comes from the work of Bernhard et al. ( 1975 ), who noted that ionic 65Zn added to some phytoplankton culture media did not exchange with the stable Zn bound to naturally occurring organic ligands, even after 20 days. More recently, Kuwabara et al. (1986) presented experimental evidence that TiO2 particles could behave similarly to dissolved chelators in controlling the free zinc concentration [Zn 2÷ ] available to algal cultures. The results of their studies imply that zinc adsorbed to TiO2 can become available to the algae at some point during the culturing period. The rate of desorption of zinc from TiO2 was not obtainable from those experiments, but the authors did point out that such a kinetic term may be essential for the accurate prediction of algal response in natural waters. Furthermore, the TiO2 particles represent an appropriate model system only for environments in which bare oxide and clay surfaces constitute the main adsorption sites for zinc. In this work, a chelating ion exchange procedure based on total exchange of the free metal was used to establish the rate laws describing the release of zinc from the particulate and dissolved organic forms naturally occurring in Narragansett Bay water. In some samples, the partitioning of zinc obtained from the ion exchange experiment was also compared with independent measurements of particulate-bound zinc as well as electroactive zinc at pH 8.0, 7.0 and 6.0. METHODS

Sample collection and treatment Samples were collected into a 4-1 high-density polyethylene jug, ~ 200 m from the shoreline. After a thorough rinse with subsurface water the bottle was immersed ~ 60 cm below the surface. It was held at that depth while the researcher was swimming with fins, and was then uncapped for filling. After

ZINC IN NARRAGANSETT BAY WATER

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the bottle was filled, it was capped underwater to avoid contamination from the surface microlayer. The sample was then brought to the laboratory where a 1-1 portion was immediately filtered for use in the last step of the resin conditioning procedure (see below).

Electrometric measurements The salinity of all Narragansett Bay water samples was determined in the laboratory using an in situ inductive salinometer (Yellow Springs Instruments Co., model 33 ). The pH measurements were made with a Coming 130 pH meter and an Orion combination Ross electrode calibrated with NBS buffers. Measurements by anodic stripping voltammetry (ASV) were made with a rotating disk electrode (Beckman) connected to a PAR model 174A polarographic analyser which was interfaced to a PAR model 315 controller (EG&G Princeton Applied Research); pH control in the electrochemical cell was achieved by controlling Pco~ of the N2/CO2 gas. For a dynamic system such as Narragansett Bay, the perception of how much metal is tied up in electro-inactive forms may depend on the storage period before analysis (Muller and Kester, 1991 ). This is why the ASV measurements were undertaken within 2 h of sampling, whereas the chelating batch experiments were usually started on the following day. The characteristic 'time of measurement' of ASV is much shorter than what can be achieved by the chelating batch method (Muller and Kester, 1990); it follows that the forms of zinc detectable by ASV should appear 100% labile to the chelating batch method, and hence the amount of zinc taken up by the resin after 0.5 h (designated Zn 1 in Table 2, below) should never be less than the amount electroactive at natural pH. This condition of internal consistency was met in every experiment. As a crude characterization of the acid-base properties of the mixture of ligands responsible for the complexation of zinc, voltammetric measurements were also undertaken at pH 7.0 and 6.0.

Chelating batch experiments The utility of the Chelex-100 chelating resin for distinguishing between different forms of a trace metal was suggested from several earlier works (Florence and Barley, 1976; Figura and McDuffie, 1980), and the technique and sampling procedures used in this work have been described in detail elsewhere (Muller and Kester, 1990 ). Briefly, a 1-1 sample of bay water is reacted continuously with 10 g of Chelex- 100 resin for up to 150. At selected time intervals, the stirring is momentarily stopped to allow the resin to settle and to sample the reactor. The trace metals in the subsamples are then preconcentrated by coprecipitation with cobalt and ammonium pyrrolidine dithiocarbamate (APDC), using filtration to collect the precipitate. The precipitate is

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dissolved in nitric acid and analysed for zinc and cadmium by atomic absorption spectrophotometry (AAS). The amount of unreacted metal can thus be measured with time. Labile metal - a large fraction of which is in the form of various inorganic complexes - corresponds to the amount of metal removed during the initial 0.5 h of the experiment. After this initial rapid removal, the more inert forms of the metal begin to dissociate in response to the extremely small pool of free metal supported by the resin. Provided the characteristic removal rates of these forms are sufficiently far apart from each other, concentrations and dissociation rate constants can be calculated from the overall removal curve. Before its introduction into the reactor (time zero), the resin is sequentially equilibrated with several filtered portions of the sample to be analysed; this procedure helps maintain a constant pH ( _+0.15 pH unit) during the time the sample is reacted. One advantage of the method is that both dissolved and particulate forms of the metal can be measured from the same experiment. This procedure provides an alternative to more ambiguous methods that involve filtration followed by chemical extraction of the metal from different binding sites (Tessier et al., 1979; Slavek et al., 1982; Trefry and Metz, 1984). However, a separation step such as filtration through a 0.40-/zm Nuclepore membrane filter will ultimately be required for an approximate distinction between dissolved and particulate fractions in the sample. Another advantage resides in the mechanistic - as opposed to empirical - approach: the species do not have to conform to some arbitrary classification; instead, they define themselves as kinetic signals against a background of more inert species. The main limitations of the method are twofold: (1) although representative subsamples containing low-density, permanently suspended particles, can easily be obtained, it is likely that larger, denser particles originating from the resuspension of sediments will be missed; this will no doubt occur if their settling velocity is comparable with that of the Chelex resin beads; (2) metal-organic complexes with stability constants less than l06 mol-1 kg may be directly taken up by the resin as a result of competition between the organic ligands in solution and the iminodiacetic groups on the resin (Cox et al., 1984); when this occurs, the metal-organic complex will be mistakenly assigned to the labile fraction. The design of Experiment One (November 1982) differed from that of later experiments in that a series of 250-ml polyethylene bottles, each corresponding to a given reaction time, were used to achieve the desired range of contact times. Each bottle contained a 1-g portion of the treated resin. A 100g subsample of either filtered or unfiltered bay water was added to each bottle and the bottle was continuously agitated with a mechanical shaker. At designated times - which ranged from seconds to tens of hours - the bottles were taken and the resin was separated out with a Teflon cloth (100-#m mesh opening, Savillex Corporation), the solution phase being saved for acid diges-

ZINCIN NARRAGANSETTBAYWATER

17 5

tion. The solutions were acidified with 0.5 ml of a mixture made up of 70% HNO3 and 70% HC1 (J.T. Baker, Ultrex), then ultrasonicated in a hot water bath (90 °C ) for 20 min to dissolve any particulate forms of zinc. After allowing the samples to cool, 10 ml of 2 M NH4OAc buffer prepared from Ultrex HOAc (glacial) and Ultrex NH4OH (20% as NH3) was added, to bring the pH to ~ 5.3. The samples were then analysed for total Zn by ASV. RESULTS

Correlation of zinc with salinity in Narragansett Bay The bay water samples that provided the source of these data were collected at either of two sampling sites located 200 m from the shoreline in the upper and lower bay respectively (Fig. 1 ). Most of them were taken in the spring and autumn of 1985 and 1986, regardless of the state of the tide; as a result, the Zn vs. S % plots of Fig. 2 include composite variations in the Zn vs. S%0 relationship within the bay and over time. Considering the variability of

Fig. 1. Map of Narragansett Bay, showing the two stations sampled (full squares).

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80

(a)

o o

Total Zn n

60

0 0

E E (D Z

0

0

O

40 0

t3

O O

0

20

El

0

n

I

80

(b) "5

E

Dissolved Zn

60 0

o ooa

40

0

0

n O 0

20

n

0

[]

i ¸

i

0

30

(c) P a r t i c u l a t e Zn 0 v

20

0

Y:

O

0

[]

u~

0

10

[]

O~
0.25

Cd3

-

k3

.--labile C a d m i u m inert--,

0.05 0.03

Cd4

5.4 7.1 8.7 6.7 5.7 14.9 74.6

Zn

0.05 0.03 0.00 0.00 0.00 0.30

Cd

Particulate fraction

Concentrations are expressed in n m o l kg - 1. Subscripts denote the degree o f lability o f the Zn and Cd forms: 1 = labile, 2 = moderately labile ( d i s s o l v e d ) , 3 = slowly labile (particulate), 4 = inert. Rate constants are in h - ~ (k2; also k3, italic value) or in n m o l k g - ~ h - ' (k3). H y p h e n s indicate values that were below the detection limit. A sample is f r o m the u p p e r bay where S < 28.5%o, a n d f r o m the lower bay otherwise.

29.7 29.2 29.4 29.0 29.3 27.4 20.7

Nov. 9, 82 May 5,85 June 3, 85 Sep. 14, 85 Oct. 10,85 Nov. 23, 85 Jan. 29, 86

(~)

S

Date

Analytical results from Narragansett Bay

TABLE 2

OO

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lution; 'moderately labile zinc' might denote a first-order dissociating organic complex present in very variable concentrations and with a characteristic kd= 0.36_ 0.08 h-1; 'slowly labile zinc' encompasses virtually all of the particulate forms of zinc retained on a 0.40-/~m filter, and its response to a reduced ambient free metal concentration is a linear release of the zinc with time, at a rate of O. 12 _ 0.04 nM h - 1. DISCUSSION

From the results presented above, it is obvious that the distribution of zinc among the kinetically defined fractions Zn 1-Zn4 is highly variable with sampling date, and perhaps state of the tide. Nevertheless, little variation is observed in the values of the kinetic rate constants k2 and k3 corresponding to the moderately labile and slowly labile fractions respectively. Indeed, results from our chelating batch experiments with Narragansett Bay water suggest the existence of characteristic reaction times for the release of zinc from the organic and/or particulate associations that occur in a given environment. Hence, the k2 and k3 values determined here might be appropriately used in any description of zinc cycling in Narragansett Bay that would incorporate kinetic terms. The kinetic signal contributed by the particulate forms of zinc is a domiaant feature in all the samples examined. However, the actual transport and reaction mechanisms that provide the sequence of paths which ultimately accounts for the observed rate is far from obvious. Experiment Seven may be seen as an exception in this regard, because a simple mechanism consisting of a single desorption step (apart from convection in the bulk sample and film diffusion at the resin bead surface, both of which are unmeasurably fast in the context of the experiment) is capable of accounting for the observed firstorder rate. In contrast, Experiments One-Six exhibit zero-order kinetics which cannot be explained simply. Perhaps the results ought to be interpreted within the context of biological processes. For instance, the work of Shannon and Lee ( 1966 ) indicates that the rate law for the hydrolysis of tripolyphosphate can be altered dramatically from first order under aseptic conditions to zero order when micro-organisms are present. As we did not endeavour to operate under sterile conditions, we cannot exclude the possibility that some of the rate-determining reactions might have been catalysed by micro-organisms. Another possibility is that the overall kinetics of release from particles is governed by a multi-stage process in which zinc first migrates rapidly from a site within the particle (biological cell?). The second stage would involve complexation at the particle surface, and the third stage would be the desorption of the metal from the surface sites. This interpretation would lead to the observed zero-order kinetics in situations where the last stage was sufficiently slow to be rate determining.

ZINC IN NARRAGANSETT BAY WATER

185

Whatever the exact nature of the zinc-particle interactions may be, it is interesting to note that under the limiting conditions that form the basis of the chelating batch technique, most of the metal can be released in less than 3 days. In natural systems, such interactions may result in a zinc-buffering action over comparable or longer time-scales, whenever [Zn 2+ ] decreases through metal uptake. For uptake processes with time-scales on the order of weeks or longer, a kinetic description of the release process is no longer required and the buffering capacity of the system can be estimated from equilibration computations in the manner described by Hirose and Sugimura (1985). Finally, it is interesting to note (Table 2 ) that the chelating batch estimates of'labile zinc' do not correlate in a simple manner with the ASV estimates of 'labile zinc' at pH 6, 7 or 8. In other words, the ASV technique cannot be calibrated against the chelating batch technique by simply adjusting the pH of measurements. This result is not surprising in view of the known chemical non-specificity of the ASV measurement in seawater (Florence, 1986). Consequently, a chemically blind approach to speciation, such as the one combining ASV at pH 6.3 and chelation in a stirred reactor for 3 days (Figura and McDuffie, 1980), does not permits comparisons of speciation other than between samples of similar composition. CONCLUSIONS

Our aim has been to characterize the various forms of zinc present in Narragansett Bay water, based on the differing rates at which Zn 2+ (or, more accurately, Chelex-reactive zinc) can be released from those forms. The dissolved forms of zinc which are not directly available for uptake by the Chelex100 resin seem to follow the expected first-order dissociation kinetics. In contrast, no straightforward and plausible explanation can be offered for the zeroorder release of Zn 2+ from suspended particulate matter in samples Nos. 1-6. Furthermore, such a rate law cannot be easily incorporated into a kinetic model which would include the rates of both chemical reactions and transport processes (Hoffmann, 1981; Morgan and Stone, 1985) relevant to Narragansett Bay (Pilson, 1985). In sample No. 7, the nature of particulate zinc is better understood, and conclusive evidence exists, that suggests desorption of zinc from organic particles or coatings; nevertheless, we do not have as yet direct observations for the kinetics of the reverse reaction (s) under the same set of conditions, and this limits the applicability of the batch uptake results presented here. Until methods are developed for the identification of specific metal-binding sites on or within natural particles, it might be useful to complement the chelating hatch measurements with an independent characterization of the metal-particle associations, as produced by extraction procedures (Tessier et

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al., 1979). Another line of approach would be to gather kinetic data on pure solid phases (metal oxides, calcium carbonates) or particles of similar composition and morphology (humic aggregates, phytoplankton cultures). REFERENCES Bernhard, M.E., Goldberg, E.D. and Piro, A., 1975. Zinc in seawater, an overview. In: E.D. Goldberg (Editor), The Nature of Seawater. Dahlem Konferenzen, Berlin, pp. 43-68. Cox, J.A., Slonawska, K. and Gatchell, D.K., 1984. Metal speciation by Donnan dialysis. Anal. Chem., 56: 650-653. Figura, P. and McDuffie, B., 1980. Determination of labilities of soluble trace metal species in aqueous environmental samples by anomic stripping voltammetry and Chelex column and batch methods. Anal. Chem., 52: 1433-1439. Florence, T.M., 1986. Electrochemcial approaches to trace element speciation in waters - a review. Analyst, 111: 489-505. Florence, T.M. and Batley, G.E., 1976. Removal of trace metals from sea water by a chelating resin. Talanta, 23: 179-186. Hirose, K. and Sugimura, Y., 1985. Role of metal-organic complexes in the marine environment. Mar. Chem., 16: 239-247. Hoffmann, M.R., 1981. Thermodynamic, kinetic, and extrathermodynamic considerations in the development of equilibrium models for aquatic systems. Environ. Sci. Technol., 15: 345353. Kuwabara, J.S., Davis, J.A. and Chang, C.C.Y., 1986. Algal growth response to particle-bound orthophosphate and zinc. Limnol. Oceanogr., 31:503-511. Morgan, J.J. and Stone, A.T., 1985. Kinetics of chemical processes of importance in lacustrine environments. In: W. Stumm (Editor), Chemical Processes in Lakes. Wiley, New York, pp. 389-426. Muller, F.L.L. and Kester, D.R., 1990. A kinetic approach to trace metal complexation in seawater: application to zinc and cadmium. Environ. Sci. Technol., 24: 234-242. Muller, F.L.L. and Kester, D.R., 1991. Voltammetric determination of the complexation parameters of zinc in marine and estuarine waters. Mar. Chem., 33:71-90. Pilson, M.E.Q., 1985. On the residence time of water in Narragansett Bay. Estuaries, 8: 2-14. Shannon, J.E. and Lee, G.F., 1966. Hydrolysis of condensed phosphates in natural waters. Air Water Pollut. Int. J., 10: 735-756. Slavek, J.E., Wold, J. and Picketing, W.F., 1982. Selective extraction of metal ions associated with humic acids. Talanta, 29: 743-749. Tessier, A., Campbell, P.G.C. and Bisson, M., 1979. Sequential extraction of dissolved components in freshwater and seawater at 25 °C and l atm pressure. Geochim. Cosmochim. Acta, 45: 855-881. Trefry, J.H. and Metz, S., 1984. Selective leaching of trace metals from sediments as a function ofpH. Anal. Chem., 56: 745-749.